The Effect of Applying Model Nanoplastic Particles to Soil on the Composition of Its Microbial Community

Abstract

Soil microorganisms play pivotal roles in biogeochemical cycling and plant growth promotion, directly impacting crop productivity and ecosystem stability. While assessing their responses to emerging contaminants like micro/NPs is critically important, research remains challenging due to highly variable effects contingent upon (1) soil physicochemical properties and (2) plastic characteristics (type, size, morphology, concentration, and other parameters). A one-month laboratory incubation experiment using 0.55 µm polystyrene latex nanoplastics (NPs) allowed us to investigate the microbial communities in soils in the southern taiga zone (near Saint Petersburg, Russia) react to the addition of polystyrene NPs. It was found that sandy Podzols were more resistant to the addition of NPs than loamy Retisols and Fluvisols. The most responsive components of the soil microbiome were those that were initially more abundant. These include representatives of the following phyla: Pseudomonadota, Bacillota, Actinomycetota and Planctomycetota. The alpha diversity parameters of the microbial community, expressed in the number of operational taxonomic units and bio-diversity indices (Shannon and Simpson), decreased under the influence of NPs. The dynamics of alpha diversity of the microbial community were the least pronounced in Podzol soil. Beta-diversity parameters changed the most in Hortic Retisol, slightly less in Fluvisol, and not at all in Podzol. Thus, it was found that agricultural soils were most affected by NPs (in terms of microbial community dynamics) compared to the region’s two zonal soils. Studies carried out indicate that, in the future, a threshold for the harmfulness of NPs in relation to soils should be developed, taking into account the differentiation of soils as standardized objects in terms of particle size distribution.

1. Introduction

Plastics have been actively used in industry and everyday life since the 1940s and 1950s. However, recent decades have witnessed a significant increase in plastic production, much of which remains uncontrolled, leading to substantial environmental pollution. Annual global plastic production increases year by year, reaching 400.1 million tons in 2020 and projected to reach 33 billion tons by 2050 [1]. The breakdown and weathering of large plastic debris lead to the formation of microplastics (MPs) and nanoplastics (NPs): insoluble, synthetic solid particles or polymeric matrices measuring less than 5 mm (MPs) and 1–1000 nm (NPs) that are not easily degraded [2]. As well as ‘secondary’ MPs originating from plastic waste products, various types of MPs are also used in cosmetics, industrial abrasives, agriculture and many other industries. MPs typically consist of polymers such as polyethylene (PE), polyester (PES) and polyamide, and often contain chemical additives such as plasticizers, flame retardants and UV stabilizers, which can leach into the environment [3]. Due to their small size and persistence, MPs are widespread pollutants found in various ecosystems and pose risks to the environment and human health.

Despite growing awareness of microplastic pollution as a worldwide environmental threat, soil pollution caused by plastics continues to be overlooked. However, the soil environment is considerably affected by the input and accumulation of MPs [1]. The main sources of MPs in soil are sewage sludge management, composting, plastic film mulching, airborne deposition and the use of agricultural machinery [2]. Recycling organic waste by composting and applying it as fertilizer on agricultural land is an environmentally friendly measure to return trace elements, nutrients and humus to soils. But most of the livestock and poultry manure is contaminated with MPs. Possible sources of these MPs include the following: extensive use of plastic products in poultry and livestock production (feed packaging bags, feed transportation pipelines, etc.); ingestion of MPs contaminated feed by livestock and poultry, which subsequently excrete MPs in their feces; MPs from the environment can get into manure during composting and transportation. Soil amendments and irrigation are considered the most important sources of MPs in agricultural soils, potentially doubling or tripling the number of MPs compared to undisturbed areas [3,4]. MPs are persistent in soils; long-term field studies have demonstrated their stability [5]. Their spatial distribution depends on soil properties such as void ratio and hydraulic conductivity, which influence their vertical and horizontal migration within the soil profile.

The adverse effects of MPs and NPs on the physical, chemical and biological properties of soil have been observed previously. MPs can alter the physicochemical characteristics of soil, such as its bulk density and water holding capacity [6]. However, some of the observed changes may be due to the lower density of plastics rather than increased porosity [7]. Furthermore, MPs can release harmful chemicals, such as phthalates and chlorinated compounds, into the soil, where they can seep into groundwater [8]. Their surfaces can also carry pathogens, acting as vectors for disease in soil environments. Previous research has shown that soil fauna such as earthworms, mites and nematodes ingest and transfer MPs, which can disrupt their health [9]. Another field experiment found that the addition of MPs and NPs significantly reduced the abundance of, and altered the community composition of, soil fauna such as microarthropods and nematodes, and that these changes cascaded through the soil food web, affecting soil ecosystem functioning.

The presence of MPs and NPs in the soil environment can considerably affect the properties of the soil microbiota, which could pose a threat to the stability of terrestrial ecosystems. One of the most serious aspects is the change in community structure that directly affects soil biodiversity. MPs and nanoplastics create a new habitat for microorganisms, known as the ‘plastisphere’, which may harbor specialized communities that are distinct from the native soil microbiota. Recent findings demonstrate that certain types of plastic, such as polypropylene and expanded polystyrene, selectively enrich specific taxa, including microorganisms that degrade plastic, which stimulate CO₂ emissions from agricultural soils [10]. Certain bacterial groups, such as Patescibacteria, have been found to be closely associated with microplastic contamination, indicating that specific microbial taxa may thrive in soils affected by MPs [11]. Furthermore, the presence of MPs can increase the relative abundance of nitrogen-fixing and phosphorus-solubilizing bacteria while reducing populations of nitrifiers and ammonia-oxidizing bacteria, thereby altering the dynamics of nutrient cycling in soils [12].

Although some studies report no significant change in overall microbial diversity, microplastics can reduce the complexity and stability of microbial networks [13]. This includes a decrease in network size and connectivity, suggesting that microplastics may favor deterministic over stochastic processes in community assembly. The introduction of microplastics influences the composition and stability of soil microbial networks. Microplastics support the proliferation of Bacteroidota and Chloroflexota, which break down the plastic, but this process reduces overall microbial richness and network complexity [14]. Furthermore, microplastics can contribute to the increased abundance of pathogenic microorganisms, potentially disrupting the stability of microbial networks and increasing the system’s sensitivity, particularly in fertilized soils.

MPs and NPs can have varying effects on soil microbial diversity depending on both the MPs type, concentration, size, and polymer composition, and on soil properties, including texture. For example, at a concentration of 5% (w/w), larger polyethylene (PE) particles (150 μm) in a clay soil reduced the richness and diversity of both bacterial and fungal communities, while smaller particles (13 μm) enriched these populations [15]. The addition of 5% polyvinyl chloride (PVC) and 1% and 5% PE to an acidic loam soil decreased soil alpha diversity; however, 1% PVC had no effect on soil bacterial diversity [16].

Soil microorganisms can adapt to the presence of microplastics by forming biofilms and utilizing MPs as substrates for biodegradation. This process involves the secretion of enzymes that break down MPs into smaller, more manageable molecules. A recent study showed that microplastics significantly reduce the activity of key soil enzymes, such as β-glucosidase, urease and dehydrogenase, by approximately 32%, 40% and 50%, respectively, with greater reductions occurring at higher concentrations of MPs. This decline correlates with a decrease in bacterial populations and strains, indicating impaired microbial function. The impact on enzymatic activities depends on the shape of the MPs and the type of polymer. For instance, polyethylene (PE) and polyvinyl chloride (PVC) MPs can increase the activity of certain enzymes, such as urease and acid phosphatase, whereas polypropylene (PP), PES and PVC can either inhibit or enhance fluorescein diacetate hydrolase activity, depending on the type of polymer [17]. The shape and type of MPs also have a different impact on community structure. For instance, PVC treatments showed higher relative abundances of families such as Acholeplasmataceae and Lachnospiraceae compared to other petroleum-based polymers such as polyethylene (PE) and polyurethane foam (PUF) [18]. Additionally, biodegradable polymers such as polylactic acid (PLA) tend to increase bacterial diversity compared to conventional plastics, stimulating the growth of certain groups such as Bacillota and Desulfobacteriota [19]. However, the addition of MPs, especially PLA, weakens microbial interactions and community stability. Among different shapes, fibers and foams significantly alter bacterial composition more than particles, with community dissimilarity increasing with microplastic concentration [12]. In a 2020 microcosm experiment using salt marsh sediment amended with polyethylene (PE), polyvinyl chloride (PVC), polyurethane foam (PUF), or polylactic acid (PLA) microplastics, it was demonstrated that bacterial diversity (alpha-diversity metrics) was highest in the biopolymer (PLA) treatment and lowest in PE-amended sediments. Across all samples, taxa from the phyla Bacteroidota and Pseudomonadota dominated. Within Pseudomonadota, the classes Deltaproteobacteria and Gammaproteobacteria were predominant [20].

MPs can influence the physicochemical properties of soil, which in turn affect microbial activity and structure. Increased soil compaction and reduced oxygen diffusion favor facultative or obligate anaerobes such as Pseudomonadota, while reducing aerobic groups such as Actinobacteriota [1]. Acidobacteriota are generally abundant in soils but often decrease in relative abundance with microplastic pollution, possibly due to changes in soil pH and nutrient availability [1]. The toxicity arising from organic and inorganic contaminants associated with MPs, such as plastic additives and adsorbed pollutants, may further modulate microbial activity and ecosystem functions [13]. Furthermore, the presence of MPs can have cascading effects through soil food webs, impacting microbial communities and overall ecosystem functioning [10].

Since soil microorganisms regulate biogeochemical cycles and support plant growth, microplastic-induced microbial alterations may reduce soil quality, crop yields, and ecosystem resilience. However, there is a lack of studies on the influence of MPs and especially NPs pollution on the soil microbiome due to numerous limitations, including soil type and properties, land use and NPs type. The escalating levels of pollution highlight the need for a nuanced, soil-type-specific approach to contamination regulation and bioremediation in both natural and anthropogenically altered soils. Addressing this need requires experimental investigations into the differential responses of microbial communities from natural versus anthropogenically modified soils to micro- and nanoplastic pollution. The main aim of this study was therefore to estimate the response of soil microbial communities to NPs in different soil types. To achieve this, the following objectives were set: (1) to set up a model experiment on NPs application to different types of soil; (2) to isolate total DNA from experimental (microplastic-exposed) and control soils for molecular analysis (3) to characterize the taxonomic diversity of soil microorganisms using high-throughput next-generation sequencing methods.

2. Materials and Methods

2.1. General Design of Experiment

The experiment included six stages (Figure 1). The first stage is the sampling of soil from three key sites with different geochemical conditions in different landscape positions.

The second stage is sample preparation of soil samples for an incubation experiment. The third stage is the addition of plastic particles in a known concentration into soil samples (samples were also left without application for comparison). The fourth stage is the incubation of samples. The fifth stage is DNA isolation and sequencing. The sixth stage is the chemical analysis of soil samples remaining after incubation. A detailed description of the sampling site and laboratory treatment is described below.

2.2. Soil Sample Characterization and Preparation

Three soil types, which are characteristic of north-west Russia and have various landscape binding options, were chosen to test the effect of NPs contamination. M1—Soil from fallow agricultural field (Hortic Retisol)—59.890016 N 29.828190 E. M2—Marsh soil (Gleyic Fluvisol)—59.900505 N 29.838823 E. M3—Sandy podzolic soil (Entic Podzol Folic)—59.947344 N 30.714777 E. The soil profiles are presented in Figure 2. The Retisol (M1) was located on an abandoned experimental agricultural field belonging to the former Biological Research Institute of Saint Petersburg State University. The field is now covered by a wild meadow. The field is situated on the second marine terrace of the Litorina Sea, which is composed of Holocene lake sediments with a clay texture (<0.001 mm—11.76%, porosity—59.07%). The Fluvisol (M2) is located on the first marine terrace of the Gulf of Finland. The parent materials here are modern marine clays that are blue in color due to the gleyification process (<0.001 mm—8.52%, porosity—32.16%). The Entic Podzol (M3) is located on the Koltushi upland, parent materials are represented by sands and sandy loams of fluvioglacial genesis (<0.001 mm—12.19%, porosity—61.90%) [20].

The altitude at sea level for points M3, M1 and M2 is, respectively: 60, 15 and 38 m. Thus, we consider the soil to be of the automorphic upland type (M3), as well as the soils of catchment basins with accumulative geochemical conditions of plains (M1) and marine terraces (M2). The territory is located in the southern taiga (boreal coniferous forests) zone of a humid continental (hemiboreal) climate (Dfb). Winters are long but mild (the average January temperature is −10 °C) and summers are warm but short (the average July temperature is 16–17 °C), with an annual total precipitation of 550–850 mm [21].

2.3. Soil Preparation for Incubation Experiment

Soil samples were collected from the 0–10 cm layer, placed in aluminum foil bags, and transported to the laboratory. Following transportation, the samples were air-dried to a constant weight. After drying, the soil was sieved through a 1 mm mesh, and roots as well as extraneous materials were removed. The sieved samples were homogenized and divided into two portions. Plastic particles were added to the first portion for the incubation experiment, whereas the second portion was incubated without plastic particles. Thus, the experiment yielded nine non-contaminated samples (three replicates per soil type) and nine plastic-amended samples (three replicates per soil type). The incubation conditions are described below.

2.4. Microplastic Preparation and Application

Polystyrene latex with a particle diameter of 0.55 µm and a concentration of 1000 µg/mL was used as a contaminating agent. Particles of this size are classified as NPs [22]. Polystyrene (PS) was chosen as the model NPs contaminant due to its widespread use in various applications, including packaging, consumer products, and agriculture, making it a prevalent microplastic and NPs type in the environment. Polystyrene latex particles, in particular, offer the advantage of uniform size and shape, facilitating controlled laboratory experiments.

The experiment was conducted in 50 mL Screw Microtubes, Skirted, Graduated, made of high-clarity, ultra-pure polypropylene (meeting FDA 21 CFR 177.1520 and USP Class VI standards). These were closed with plastic Screw Cap Tubes also made of Ultraclear medical-grade polypropylene. The material of both the tubes and caps is certified free of detectable RNase, DNase, DNA, and PCR inhibitors. Due to the fundamental importance of this last characteristic for obtaining reliable results, we decided to use these plastic items in the experiment, which aimed to assess the influence of the presence of plastic in various soils, even though their use increased the risks of contamination by NPs from external sources. To minimize variability in results, all experiments utilized tubes and caps made of identical material, thoroughly rinsed prior to use (all laboratory glassware, including the described tubes and caps, was rinsed three times with ultrapure water (0.2 μm pore size). Furthermore, we confirmed that the added nanoplastic (polystyrene) and the material from which the tubes and caps were made (polypropylene) are different types of plastic, and that the concentration of polystyrene was many times higher than the potential contamination from NPs leaching from the walls of the laboratory vessels. This suggests that the results are primarily due to the influence of the added polystyrene.

From the previously prepared soil samples of three types, six 30 g aliquots were taken from each type, totaling 18 aliquots. These aliquots were transferred to test tubes such that for each soil type, three aliquots were amended with plastic particles and three were left unamended. Tubes were filled with 7.5 mL of plastic solution—approximately 250 μg of plastic per gram of soil. This quantity of microplastic simulated 0.025% contamination by weight. Blank samples were supplemented with 7.5 mL of ultrapure water.

NPs concentrations used in this study approximated levels found in natural settings: their concentrations show considerable variability, ranging from a few particles/kg in regions with low inputs to more than 80,000 particles/kg in areas with long-standing mulching practices. Experimental concentrations used in many studies from 0.01% (m/m) to 7% (m/m) are well above typical levels in the environment. 1% by weight corresponds to 10,000 mg MPs per kilogram of soil, which is much higher than typical MPs concentrations currently found in agricultural soils. They typically vary from a few hundred to a few thousand particles per kilogram, corresponding to much lower mass concentrations, often below 0.1% (m/m) [23].

To ensure consistent conditions across all treatments, control vessels were adjusted to a moisture content equivalent to approximately 60% of the soil’s total water-holding capacity by adding 7.5 mL of sterile distilled water. Following thorough mixing, all tubes were aerated via small holes created in the base and lid with a sterile awl. During a one-month incubation at 23 °C, the tubes were weighed weekly, and the amount of water lost to evaporation was replenished with sterile distilled water. All samples were then frozen to halt microbial activity before subsequent analyses.

2.5. Chemical Analysis of Soil Samples

Following the incubation experiment, basic chemical analyses were performed on soils without NPs addition (in nature conditions), and soil samples into which plastic particles were artificially added. Some studies emphasize that NPs can affect fundamental soil properties, such as microbiological activity, carbon content, pH and density. The pH values were measured using the potentiometric method at a soil-to-water ratio of 1:2.5 and at a soil-to-0.01 N CaCl₂ solution ratio of 1:5 [24]. Soil carbon content and total nitrogen were determined using a CHN analyzer Euro EA3000 3028-HT-OM (EuroVector Instruments & Software, Maestà, Italy). The available forms of phosphorus and potassium were determined using the Kirsanov method [25]. Cation exchange capacity (CEC) was measured in bulk soil samples using the Bobko–Askinazi–Aleshin method in a BaCl₂ buffer solution (pH = 6.5) [26].

2.6. DNA Processing

Total DNA was isolated from each tube in triplicate using the RIAM protocol [27], resulting in 54 samples of DNA in total. The Research Institute for Agricultural Microbiology (RIAM) protocol is a universal method for isolating high-purity DNA from diverse soil types. This protocol avoids silica column purification, utilizing high concentrations of phosphate buffer to prevent DNA sorption to minerals. DNA is then precipitated using CTAB. RIAM offers superior DNA yield and purity compared to commercial kits, with no observed PCR inhibition.

All samples were amplified using primers 515F (GTG CCA GCMGCC GCG GTAA) and 806R (GGA CTA CVS GGG TAT CTAAT) [28] to target the variable region of the 16S rRNA gene and were then prepared for pair-ended 2 × 300 bp sequencing on the Illumina MiSeq platform, as previously described [29,30].

2.7. Statistics

The statistical processing of the results of the soil chemical analysis was performed using one-way analysis of variance (ANOVA) with Tukey test (alpha > 0.05) in Origin Pro software (version 2024; OriginLab Corporation, Northampton, MA, USA).

The DNA data were analyzed using the DADA2 pipeline in the R software environment (v. 4.2) [28,31,32]. The truncation lengths were 220 bp for forward and 170 bp for the reverse sequences. Error rate was set to 2.4 and minimum overlap was set to default. The reads were clustered into ASVs (amplicon sequence variants), to which taxonomy was assigned using the SILVA 138.1 database. Before the formal analysis data were rarified. Alpha diversity [33,34], beta diversity [35], CCA (canonical correspondence analysis) [36] and ANCOM-BC (analysis of compositions of microbiomes with bias correction) [37] were performed using the vegan [38], phyloseq [39] and ampvis2 [40] packages to assess changes in soil microbiomes with the addition of NPs. For the ANCOM-BC output we selected ASVs, increasing (with logFC > 1) or decreasing (with logFC < 1) in contaminated soil with padj < 0.05.

3. Results and Discussion

3.1. Soil Chemical Properties

Following the incubation experiment, basic chemical analyses were performed on control soil samples and soil samples containing plastic contamination. Some studies emphasize that plastics can affect fundamental soil properties, such as microbiological activity, carbon content, pH and density [41,42,43,44,45]. Our research found virtually no effect of NPs addition on pH, total carbon and nitrogen content, available forms of potassium and phosphorus, and cation exchange capacity (Figure 3). Only statistically significant (p < 0.05) effects were found for cation exchange capacity (3.30 ± 0.04 vs. 3.77 ± 0.15 meq/100 g) in soil M1 (Hortic Retisol and for pH in the water extract (6.0 ± 0.1 vs. 5.8 ± 0.1) in soil M3 (Entic Podzols). Changes in the other parameters in all soil samples studied were not statistically significant (p > 0.05).

3.2. Soil Microbiome

These data are presented in Figure 4. Notably, the addition of NPs to soils M1 and M2 leads to a decrease in the number of OTUs. Such patterns were not observed in the background podzol, which may indicate a greater resistance of the microbiota present in sandy soils to NPs contamination. However, in the case of clayey soils (M1 and M2), the addition of NPs apparently leads to decreased soil aeration, which may result in changes to soil biological parameters. The Shannon index is widely used in ecology to assess species diversity in a community. The dynamics of alpha diversity of the microbial community are least pronounced in Podzol soil. Shannon index values decrease in M1 and M2 soils. Podzol demonstrates the preservation of Shannon index values when the soil is contaminated with NPs. Similar, albeit less pronounced, patterns are characteristic of the behavior of the Simpson index, which is a measure of species diversity that takes into account both the number of species present in a community and their relative abundance. Thus, in terms of alpha diversity metrics, the addition of NPs significantly decreased alpha diversity, according to the Shannon index, in soils M1 (Hortic Retisol) and M2 (Gleic Fluvisol). The issue of regulatory concentrations of NPs in soils will arise in the future. Studies on the buffering or resistance of soils to NPs pollution will contribute to the development of standards.

Bray–Curtis metrics are used to analyze the beta diversity of a microbial community [46]. According to the beta diversity analysis (Figure 5), the addition of NPs significantly altered the soil M2 (Gleyic Fluvisol) microbiome composition. Changes in M1 soil were also noted but were not significant according to PERMANOVA. Slightly less significant differences were characteristic of the Retisol agrosoil. Minimal differences were characteristic of the benchmark sandy Podzols, confirming the conclusions obtained in the analysis of alpha biodiversity indices. It can be concluded that, in both background and polluted soils, part of the microbiome remains in the so-called ‘core’ space without experiencing a shift, while part of the microbiome, in terms of species and group composition, can be attributed to the minor component and undergoes significant changes.

At the phylum level, few changes in the microbiome were observed with the addition of microplastics (Figure 6). Pseudomonadota was the dominant phylum in all soils, both initially and following the addition of NPs. At the same time, its content remained stable in the presence of contamination. It appears to be one of the most stable representatives of the core microbial community. The dominance and stability of this phylum under anthropogenic impacts have been noted previously for some agricultural soils. Representatives of the Bacillota phylum (formerly known as Firmicutes) constituted the second most abundant group of microorganisms in all soils studied, showing a relative increase in all soils contaminated with NPs, albeit minimal in the background Podzol. Members of this phylum are known to respond to changes in soil pH and moisture levels, both of which change when NPs are added to the soil. The increase in Bacillota may be linked to their physiology and functional role in the degradation of NPs. It was shown that plant-growth promoting strains of Bacillus are capable of oxidizing the plastic in low dosage [47]. The proportion of Actinomycetota representatives barely changed, showing a slight trend towards decreasing abundance. These microorganisms are usually confined to the upper soil horizons with a high humus content. Therefore, if we continue the experiment with lower soil layers in the future, the results may differ. Planctomycetota are typical representatives of soil microbial communities, so it is not surprising that they rank fourth in terms of occurrence at the sites we studied. These microorganisms are presumably sensitive to soil moisture, so they may be more abundant in clayey and loamy soils than in sandy soils, as we observed.

As for representatives of phyla of microorganisms with an average percentage, they did not show dynamics when adding microplastics in all soils studied. These include the following phyla of microorganisms: Bacteroidota, Verrucomicrobiota, Acidobacteriota, Myxococcota and Gemmatimonadota. At the same time, the proportion of Bacteroidota, Verrucomicrobiota, Acidobacteriota in sandy Podzol was increased compared to loamy soils—M1 and M2, although their content did not change when NPs were added to the soil. This can be considered simply a distinctive feature of the soil microbiome with different particle size distribution.

Representatives of Patescibateria, Chloroflexota, Cyanobacteria, Bdellovibrionota in all studied soils turned out to be minor groups and did not play a role in both the microbiome of the original soils and the soil microbiome with the addition of NPs.

ANCOM-BC analysis revealed that in M2 soil, the addition of microplastics/NPs resulted in a significant decrease in the representation of Rhizobium mesosinicum, Minicystis, and members of the Myxococcaceae family, a group belonging to the Myxococcota phylum. On the other hand, the representation of Tumebacillus, Bacillus gottheilii from Bacillota phylum, Massilia, Devosia, Afipia felis, Bordetella, Phreatobacter from Pseudomonadota phylum increased significantly. This coincides with the findings that members of Pseudomonadota, and Bacillota were found on plastic biofilms and showed thew ability to degrade plastics [48].

Thus, the addition of NPs to the soil only influences the abundance of phyla already dominant in the initial soil microbiome composition, causing a shift in the microbiome type. It should also be noted that only the organic matter-rich topsoil horizons were analyzed.

According to the CCA (Figure 7), changes in the composition of the microbiome correlated with changes in soil SOC. This is consistent with the previous statement that the role of the organic matter in the upper soil horizons in forming the quantitative taxonomic composition of the microbiome is significant. Organic matter can influence soil acidity and humidity, as well as acting as a substrate.

Observed changes in soil microbial community diversity following the introduction of polystyrene nanoparticles (PS-NPs) highlight significant consequences for ecosystem functioning, particularly in anthropogenically modified landscapes. Our results demonstrate that agricultural soils exposed to NPs show substantial reductions in both operational taxonomic units (OTUs) and biodiversity indices (Shannon and Simpson), while undisturbed podzolic soils exhibit greater resilience. This contrast underscores the importance of ecosystem integrity in resisting emerging contaminants, indicating the crucial role of land use history and soil management practices in shaping ecological responses to nanoplastic pollution.

These findings align with ecological theory suggesting that biodiverse, undisturbed ecosystems possess greater resilience due to niche complementarity and functional overlap among species. The implications for agroecosystems could be profound: continued accumulation of NPs in arable lands may lead to long-term effects including reduced soil fertility, disrupted nutrient cycling, and decreased agricultural productivity.

The reduction in microbial alpha diversity observed in our study corroborates emerging evidence that NPs exert selective pressure on soil microbiomes. Metagenomic analyses from recent studies indicate that even trace amounts of NPs can alter bacterial composition, promoting the growth of taxa with specific metabolic adaptations (e.g., plastic-degrading enzymes or stress-resistant phenotypes) while suppressing others. For instance, Fei et al. (2020) [16] reported increased abundance of Burkholderiaceae—a family known for nitrogen fixation and organic pollutant degradation—under microplastic stress, alongside decreased populations of Sphingomonadaceae and Xanthobacteraceae, which play key roles in xenobiotic breakdown. Such taxonomic shifts may disrupt critical ecosystem functions including organic matter decomposition, nitrogen transformation, and pollutant detoxification.

Furthermore, functional gene analysis by Fei et al. (2020) [16] revealed particular sensitivity of membrane transporters to microplastics, suggesting that cellular uptake or adsorption of plastic particles may interfere with nutrient acquisition and energy metabolism in microorganisms. Given that NPs have higher surface-area-to-volume ratios than microplastics, their interactions with microbial membranes and extracellular enzymes could be even more pronounced, exacerbating metabolic dysregulation.

The context-dependent nature of nanoplastic effects—where particle size, polymer type, and soil properties modulate outcomes—complicates risk assessment while simultaneously emphasizing the need for ecosystem-specific mitigation strategies. For example, a recent study utilizing metagenomic sequencing demonstrated that even low doses (0.01% w/w) of polystyrene NPs (nPS) significantly altered bacterial composition, highlighting the sensitivity of soil microbial communities to these contaminants [49]. In agricultural systems where soils are already stressed by intensive land use, NPs may act as an additional destabilizing factor, potentially leading to:

• Altered biogeochemical cycles: shifts in microbial communities responsible for carbon, nitrogen, and phosphorus cycling could create nutrient availability imbalances, affecting plant growth and greenhouse gas emissions.
• Reduced functional redundancy: loss of microbial diversity may weaken soil’s capacity to withstand additional disturbances like drought or pathogen outbreaks.
• Trophic cascades: as soil microorganisms form the foundation of detrital food webs, compositional changes could propagate to higher trophic levels, impacting invertebrates, fungi, and plant-microbe symbiotic relationships.

In contrast, natural ecosystems such as podzolic soils may initially buffer nanoplastic impacts due to their higher biodiversity and organic matter content, which can bind contaminants and reduce their toxicity. However, chronic exposure or synergistic effects with other stressors (e.g., climate change, acid deposition) may ultimately overcome the resilience of even these systems.

4. Conclusions

Our study experimentally demonstrates that microbial communities in natural and anthropogenically modified soils exhibit differential responses to nanoplastic contamination. The effect of NPs on podzolic and floodplain soils in the humid subzone of the southern taiga was studied in the environs of St. Petersburg. It was found that adding NPs to the topsoil has a slight effect on its chemical composition. Only two cases showed significant changes in the chemical properties of the soil after the introduction of NPs. These were cation exchange capacity for Hortic Retisol and pH in an aqueous Entic Podzol suspension. Regarding the soil microbiome, contamination with NPs can lead to a decrease in operational taxonomic units, as well as in the Shannon and Simpson indices. At the same time, loamy Retisols and floodplain Fluvisols showed greater microbiome sensitivity to nanoplastic addition than the background Podzol. A key task for future studies is to investigate how the soil microbial community responds to large concentrations of NPs. It is possible that standards for the safe content of NPs in soils will need to be differentiated according to particle size distribution and organic matter content, as well as respiratory activity of the microbial community. A comparative and detailed analysis of the effects of microplastic presence and accumulation in soils presents a significant challenge due to, on one hand, the impressive diversity of microplastic types and modifications (including those related to particle size, origin, and aging) and forms (fibers, particles, granules, etc.) that exhibit varying impacts, and on the other hand, the wide range of effects manifested across different soil ecosystems. But such studies have the potential to inform the development of a differentiated approach to contamination regulation and site-specific bioremediation strategies.